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Besides eliciting several toxic effects (immunotoxicity, hepatotoxicity, carcinogenicity),
compounds, such as PCB 77, PCB 126, Aroclor 1254 (polychlorobiphenyls), or benzo[
a
]
pyrene (BaP), can lead to reproductive process impairment in inhibiting fish VTG secre-
tion or reducing gonadal development (Smeets et al. 1999; Vaccaro et al. 2005).
Some of the most potent AhR agonists, PCBs, polychlorinated dibenzo-
p
-dioxins, or
polychlorinated dibenzofurans, can inhibit some E2 responses in rodents, suggesting a
relationship between AhR activation and antiestrogenicity. 2,3,7,8-Tetrachlorodibenzo-
p
-
dioxin can induce estrogenic action or inhibit estrogen-induced effects in various tissues
because of AhR-ER cross-talk. Inhibitory AhR and ER cross-talk have been demonstrated
in mammalian cancers (Safe et al. 1991).
Several mechanisms have been elucidated that partially explain the AhR-mediated anti-
estrogenic activities of dioxins and related compounds. These include increased estrogen
metabolism, down-regulation of ER protein levels and binding activity, and decreased
ER-mediated gene expression (Gillesby and Zacharewski 1998). The ER nuclear receptor and
AhR are both critical for respective VTG and Cyp 1A expression. Arukwe et al. (2001) stressed
that interaction between the AhR and ER in sexually maturing fish can cause adverse repro-
ductive effect through disruption of estrogen-regulated reproductive processes.
The physiological functions of the AhR during development appear to be ancestral to
adaptive functions. Consequently, differences in sensitivity to the developmental toxicity
of dioxins and related chemicals may have had their origin in the evolution of the dioxin-
binding capacity of AhR in the vertebrate lineage (Hahn 2002).
8.3.3 Effects of Complex Mixtures Containing Substances with Endocrine Potential
As underlined by Tyler and Jobling (2008), the phenomenon of estrogenic-containing
effluents has been established widely across Europe, United States, Japan, and China.
Tiny quantities of several estrogenic substances in urban effluents lead to measurable
estrogenic effects in freshwater (Jobling et al. 2002, 2003; Silva et al. 2002), and brackish
environments as shown in wild populations of flounders
Platichthys flesus
from four estu-
aries in Great Britain (Tyne, Crouch, Thames, and Mersey) (Allen et al. 1999a, 1999b). The
estrogenic effects on these fish are revealed by a 4- to 6-fold increase in VTG concentra-
tions by comparison to controls and by the presence of intersex (oocytes imbricated in
testicles). Such effects were also shown in several other fish species such as
Notropis hud-
sonius
(Aravindakshan et al. 2004),
Rutilus rutilus
(Bjerregaard et al. 2006),
Oncorhynchus
mykiss
(Gagné et al. 2006), and other salmonids (Lahnsteiner et al. 2006). Recently, Vajda
et al. (2011) linked serious perturbations, revealed as intersex, gonad deformities, and
feminization, measured in the wild white suckers (
Castotomus commersoni
) sampled in
wastewater effluent-dominated Colorado streams, to endocrine active chemicals (steroi-
dal hormones, estrogenic alkylphenols, and bisphenol A) present in the WWTP effluents
(Woodling et al. 2006).
Chemicals acting similarly as mimetic estrogens can act in combination according to
the principle of concentration addition. Consequently, when present as a mixture in the
receiving waters, they are likely to pose a significant environmental risk, even when each
substance is present at below the threshold of detectable effects (Brian et al. 2007). Such
additive effects are a topic of concern for the long-term future of exposed ichthyological
populations. Estrogenic effects have also been observed in other aquatic vertebrates such
as the frog
Xenopus laevis
(Coady et al. 2005).
Despite the coexistence of several classes of EDCs (estrogenic, androgenic) in most aquatic
ecosystems, there is still limited information regarding their combined effects. Bugel et
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